This is an Open Access article distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/2.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

Abstract

Background

Polychlorinated biphenyls [PCBs], perfluorinated compounds, and polybrominated diphenyl
ethers [PBDEs] were retrospectively analyzed in archived herring gull (Larus argentatus) eggs from the North and the Baltic Sea over the last 20 years. The aim was to assess
temporal trends and effects of regulatory measures.

Results

PCBs (sum of 7 congeners) were highest in eggs from the North Sea island Trischen,
i.e., 3,710 to 20,760 ng/g lipid weight [lw] compared to 2,530 to 11,650 ng/g lw on
the North Sea island Mellum and 4,840 to 9,190 ng/g lw on the Baltic Sea island Heuwiese.
During the study period, PCBs decreased significantly. Concentrations of PFOS ranged
between 46 and 170 ng/g wet weight [ww] at Trischen, 39 to 99 ng/g ww at Mellum, and
20 to 159 ng/g ww at Heuwiese. Since 2000 and 2003, concentration levels decreased
in eggs from Mellum and Heuwiese, respectively. Perfluorooctanoic acid was the dominant
perfluorinated carboxylic acid in the North Sea eggs (Trischen 2.0 to 74 ng/g ww;
Mellum 2.6 to 118 ng/g ww), whereas perfluoroundecanoate [PFUnA] and perfluorodecanoate
[PFDA] (means 3.9 ± 3.6 ng/g and 2.9 ± 2.3 ng/g ww, respectively) dominated in the
Baltic Sea eggs. At all three locations, longer-chained perfluorinated carboxylic
acids (perfluorononanoate, PFDA, PFUnA, perfluorododecanoate) increased during the
monitoring period. PBDE concentrations (sum of 35 congeners) in eggs were in the ranges
of 282 to 2,059 ng/g lw (Mellum), 116 to 1,722 ng/g (Trischen), and 232 to 2,021 ng/g
lw (Heuwiese). Congeners associated with commercial Penta- and Octa-BDE formulations
decreased during the study period. No decrease was observed for technical Deca-BDE.

Conclusion

Effects of regulatory measures were apparent for PCBs and Penta- and Octa-BDE, while
no consistent trend is noticeable for PFOS.

Keywords:

Background

Industrial chemicals like polychlorinated biphenyls [PCBs], perfluorinated compounds
[PFCs], and polybrominated diphenyl ethers [PBDEs] have been used in innumerable applications
for many years. In recent years, however, it became clear that these compounds are
by no means unproblematic when reaching the environment. Many of them are highly persistent,
bioaccumulative and toxic. Furthermore, the substances themselves or their precursors
are subject to long-range transport and have been detected in biota in remote areas
like the Arctic and Antarctica (e.g., [1-16]). These findings led to the inclusion of PCBs, certain PBDEs (i.e., hexabromodiphenyl
ether, heptabromodiphenyl ether, tetrabromodiphenyl ether, and pentabromo-diphenyl
ether), as well as the PFCs' perfluorooctane sulfonic acid [PFOS] and its salts and
perfluorooctane sulfonyl fluoride into the list of persistent organic pollutants [POPs]
covered by the Stockholm Convention on POPs [17,18]. Already 9 years earlier, in 2000, a voluntary phase-out of PFOS was initiated by
its main producer 3 M. Others, like perfluorooctanoic acid [PFOA] are still under
evaluation. In the European Union [EU], PCBs and technical Penta- and Octa-BDEs were
banned already in 1985 and 2004, respectively [19,20]. An EU-wide restriction of Deca-BDE followed in 2008 (curia.europa.eu; case C-14/06).

Polychlorinated biphenyls were widely used since the late 1920s, e.g., as hydraulic
fluids, lubricating oils, and cooling and insulating fluids for transformers and capacitors,
plasticizers, stabilizing additives, and flame retardants. Their unfavorable properties,
namely, persistence, lipophilicity, potential for bioaccumulation and biomagnification,
and high chronic toxicity, were recognized already in the 1930s. Some of the 209 PCB
congeners are even suspected to be carcinogenic, teratogenic, reproductively and developmentally
toxic, and endocrine-disruptive. Due to their lipophilicity, PCBs accumulate in fatty
tissues and are sequestered into bird eggs [21-23].

Perfluorinated compounds have been used since the 1950s. Their chemical properties
(e.g., heat resistance, lipophobicity, hydrophobicity) render them suitable for many
applications, including fire-fighting foams, carpet impregnation, coatings and additives,
leather and textile treatments, paper and packaging, industrial and household cleaning
products, pesticides, and insecticides [11,24]. It was only in recent years that their environmental relevance became obvious [2]. Several PFCs are highly persistent, bioaccumulative and toxic and enrich in the
food web [11,13,25-32]. Unlike lipophilic compounds like PCBs and PBDEs, PFCs bind to proteins, and concentrations
are highest in the protein fraction of the blood, liver, and kidney [32-37]. For birds, a species- and compound-specific sequestering of PFCs into eggs can be
observed [12,36-41].

Polybrominated diphenyl ethers are additive flame retardants which have been used
for more than 30 years in multiple applications, such as polymers, paints, electrical
appliances, and polyurethane foams used in furniture and cars [42,43]. Of 209 possible congeners, only about 20 to 30 are environmentally relevant as components
or degradation products of the commercial formulations Penta-BDE, Octa-BDE, and Deca-BDE
[42]. PBDEs are persistent, hydrophobic, bioaccumulative, and toxic. Some congeners act
as endocrine disruptors, and the main component of technical Deca-BDE, BDE-209, has
been classified as a possible human carcinogen [44-46]. Like PCBs, PBDEs accumulate in fatty tissues. They biomagnify in the food chain
and are effectively transferred into bird eggs [43,47-49].

The dynamics of these substances in the environment and the effectiveness of regulatory
measures can be evaluated by assessing temporal trends in biota. In this context,
the German Environmental Specimen Bank [ESB] (http://www.umweltprobenbank.dewebcite) [50-53] offers the opportunity for retrospective monitoring using archived specimens of marine,
fresh water, and terrestrial ecosystems sampled regularly since the 1980s. Within
the ESB, eggs of the herring gull (Larus argentatus) are used as specimen representing marine top predators [53]. These samples are of special interest when dealing with bioaccumulating and biomagnifying
compounds.

The aim of the present study was to evaluate temporal trends of PCBs, PFCs, and PBDEs
in herring gull eggs from islands in the North Sea and the Baltic Sea. It complements
an earlier study dealing with PFCs in aquatic organisms and bird eggs [54]. For this purpose, archived homogenate samples from the German ESB covering the periods
from 1988 to 2008 (North Sea) and 1991/1996 to 2008 (Baltic Sea) were analyzed for
PCBs (7 congeners), PBDEs (35 congeners), and PFCs (twelve individual compounds).

Materials and methods

ESB sample treatment

In order to ensure a high degree of continuity, all steps of the ESB sample treatment
are performed according to ESB standard operation procedures [50,55,56]. Herring gull eggs are collected annually from the North Sea islands Trischen and
Mellum and from the Baltic Sea island Heuwiese (Figure 1). Sampling is performed by trained personnel to minimize the disturbance of the colony.
A detailed description of the sampling procedure is given in the corresponding standard
operating procedure which can be retrieved from the website of the ESB (http://www.umweltprobenbank.dewebcite). Every year, at least 35 eggs from each site are pooled and homogenized. For this
study, a total of 87 egg homogenates were analyzed. In general, samples were analyzed
once. For determining the analytical precision, representative samples were analyzed
as replicates.

Figure 1.German ESB sampling sites for herring gull eggs in the North Sea and Baltic Sea.

Analytical methods

PCB

The analysis of herring gull eggs was performed by high-resolution gas chromatography/high-resolution
mass spectrometry [HRGC/HRMS] using isotope dilution quantification. The method followed
the mutually agreed procedure of the ESB [50] and the laboratory. Details of the analytical method and instrumentation are given
in Additional file 1.

Quality assurance measures

A number of internal and external quality assurance/quality control [QA/QC] procedures
were performed: For each batch of 8 to 10 samples, a QC control sample and a blank
sample were analyzed in parallel. Additionally, a number of samples were analyzed
in duplicate. Regarding external controls, the laboratory participated in at least
two international interlaboratory control studies per year for PCBs in biological
samples. The successful participation in the Norwegian Institute of Public Health
ring tests (Round 11, 2010; Round 12, 2011) and Quality Assurance of Information for
Marine Environmental Monitoring in Europe [QUASIMEME] (Round 62, 2010) resulted in
single analyte z-scores of < 2 in 95% of the cases (n = 91).

PFC

Analysis of herring gull eggs was performed by high performance liquid chromatography-mass
spectrometry-mass spectrometry [HPLC-MS-MS] after fat removal by n-hexane extraction of alkaline sample suspensions followed by analyte extraction from
the remaining and acidified aqueous phases using methyl tert-butyl ether [MTBE]. Details on the sample preparation, analytical methods, instruments,
and quality assurance measures are given in Additional file 1.

Validation and quality assurance measures of analytical method

The validation of the analytical method was performed by standard addition experiments.
This was because no PFC-free egg matrix was available. The recoveries of the analytes
were determined by means of the slope of the recovery functions. The respective mean
recoveries for eight different concentration levels were the following: PFBA 74%,
PFBS 112%, PFHxA 108%, PFHxS 106%, PFHpS 103%, PFOA 95%, PFNA 99%, PFOS 95%, PFDA
100%, PFUnA 101%, PFDS 76%, and PFDoA 112%. Reported data were not corrected for recoveries.
In addition the parameters specificity, linearity, working range, accuracy, precision,
and limit of quantitation [LOQ] were considered. The respective data are summarized
in Additional file 1.

PBDE

PBDE were determined by HRGC/HRMS using isotope dilution quantification. The method
followed the mutually agreed procedure of the ESB [50] and the laboratory. Details of the analytical method and instrumentation are included
in Additional file 1.

Quality assurance measures

A number of internal and external QA/QC procedures were performed: For each batch
of 6 to 10 samples, a QC control pool sample and a blank sample were analyzed in parallel.
Additionally, a number of samples were analyzed in duplicate. Regarding external controls,
the laboratory participated successfully in at least two international interlaboratory
control studies per year for PBDEs in biological samples. The successful participation
in the Norwegian Institute of Public Health ring tests (Round 11, 2010; Round 12,
2011) and QUASIMEME (Round 62, 2010) resulted in single analyte z-scores of < 2 in 88% of the cases (n = 84).

Statistics

Time trends were analyzed using a statistical program (PIA) retrieved from the Arctic
Monitoring and Assessment Programme homepage [57]. The program uses a robust regression-based analysis to assess time series [58]. The percent-annual concentration change is derived from the slope of a log-linear
regression curve. The trend analysis was based on the curves of the mean concentrations
per year and was performed for the whole study period as well as for certain periods
which obviously displayed trends. The results and relevant statistical parameters
are summarized in Table S7 in Additional file 1.

Results and discussion

During the whole monitoring period, concentrations of PCBs were clearly higher than
levels of PFCs and PBDEs in herring gull eggs from all three locations. Nevertheless,
as Figure 2 shows exemplarily for three time spans, the relative importance of PCBs has declined
significantly since the study began, while PFCs are steadily increasing at two locations
(North Sea/Trischen and Baltic Sea/Heuwiese). Concentrations of PBDEs were mostly
lower than those of PFCs and decreased during the monitoring period.

PCB

PCB contamination was highest in herring gull eggs from the North Sea island Trischen.
During the study period from 1988 to 2008, the total PCB levels (sum of 7 congeners)
ranged between 3,710 and 20,760 ng/g lipid weight [lw] (corresponding to 380 to 1,760
ng/g wet weight [ww]) with the highest concentrations between 1988 and 1992 (Figure
3) [see Table S4 in Additional file 1]. Eggs from Mellum had somewhat lower levels ranging between 2,530 and 11,650 ng/g
lw (230 to 1,008 ng/g ww). Again, contaminations were highest in the late 1980s. At
both North Sea sites, egg concentrations decreased significantly (p < 0.001) during the study period which is probably related to the EU-wide ban of PCBs
in 1985 [see Table S7 in Additional file 1]. At the Baltic Sea location Heuwiese, PCB analysis started in 1996. Total PCB concentrations
ranged between 4,840 and 9,190 ng/g lw (corresponding to 460 and 900 ng/g ww) and
decreased significantly (p < 0.038) during the whole study period. (Figure 3) [see Tables S4 and S7 in Additional file 1].

Congener patterns were similar at all three locations with CB-153 > CB-138 > CB-180
> CB-118 [see Table S4 in Additional file 1]. Lower chlorinated congeners, i.e., CB-28, -53, and -101 accounted for only 0.2%
to 6% of the total PCBs. These findings reflect the enhanced bioaccumulation potential
of higher chlorinated PCBs - especially the hexachlorinated congeners CB-153 and CB-138
- as compared to the more volatile lower chlorinated congeners [23,59].

Comparison with other monitoring data

Our findings regarding the predominance of CB-153 confirm the results of numerous
other studies (e.g., [5,60-69]). CB-153 is therefore often used as a marker for PCBs. Regarding Baltic Sea locations,
CB-153 concentrations of 2,500 and 2,300 ng/g lw are reported for guillemot eggs (Uria aalge) from the Swedish island Stora Karlsö sampled in 2005 and 2006, respectively [63,65]. These data correspond well to the levels we detected in herring gull eggs from Heuwiese,
i.e., 2,650 and 2,670 ng CB-153/g lw in 2005 and 2006.

For the North Sea, no comparative CB-153 data are available, but several studies report
on CB-153 in eggs from different North Atlantic locations. At the Faroe Island and
Iceland, CB-153 concentrations in Northern fulmar (Fulmaris glacialis) eggs ranged between 1,200 and 7,500 ng/g lw in 2001 and 2002 [61,69]. Comparable concentrations were detected in herring gull eggs from three locations
in northern Norway in 2003, i.e., 2,700 to 3,300 ng CB-153/g lw [64], as well as in eggs of the European shag (Phalacrocorax aristotelis) from the Norwegian island Sklinna in 2003 and 2004 (1,640 to 1,810 ng/g lw) [67], and in eggs of the great black-backed gull (Larus marinus, 420 to 3,400 ng/g lw) and lesser black-backed gull (Larus fucus, 770 to 3,000 ng/g lw) from Iceland in 2003 [69]. All these data are in good accordance with our findings: In the same time span (2001
to 2004), CB-153 ranged between 1,290 and 3,550 ng/g lw in herring gull eggs from
Mellum and between 2,300 and 4,300 ng/g lw in eggs from Trischen. Much lower CB-153
levels, i.e., 250 to 400 ng/g lw, were reported for guillemot eggs from different
sites in Iceland, Faroe Islands, and Norway between 2002 and 2005 [64,65,69] and from eggs of the common eider (Somateria mollissima) from Iceland and three Norwegian sites (37 to 450 ng/g lw) in 2003 and 2004 [67,69]. These lower values reflect not only different environmental contamination, but also
species-specific differences in feeding habits and metabolisms as reported by Jörundsdóttir
et al. [69] for seven seabird species from Iceland where metabolite patterns and CB-153 concentrations
varied strongly between species (i.e., 150 to 18,000 ng CB-153/g lw).

Regarding total PCB concentrations, Bustnes et al. [70] report mean levels (sum of 26 congeners) of 1,240 ng/g ww (≈12,580 ng/g lw calculated
with reported lipid concentrations of 9.85%) in eggs of the lesser black-backed gull
(Larus fuscus intermediatus) collected at the southern Norwegian coast (North Sea) in 2002. In lesser black-backed
gulls from Iceland, total PCBs (sum of 12 congeners) amounted to 6,000 ng/g lw in
2003 [69]. High median concentrations are reported for eggs of ivory gulls (Pagophila eburnea) from the Russian and Norwegian Arctic regions (sum of 28 congeners 16,000 to 45,500
ng/g lw) in 2006 [71] and for eggs of the peregrine falcon (Falco peregrinus tundris) from South Greenland (sum of 22 congeners 55 μg/g lw = 55,000 ng/g), sampled between
1986 and 2003 [72]. Both species are top predators and thus strongly exposed to biomagnifying substances
like PCBs. Shag eggs from three differently exposed islands of central Norway had
total PCB levels (sum of 12 congeners) of 4,590 to 4,690 ng/g lw in 2003 and 2004,
whereas significantly lower levels of 109 to 880 ng/g lw were detected in eggs of
the common eider from the same sites [67]. These findings were ascribed to differences in feeding habits and ecology. In eggs
of the Eurasian oystercatcher (Haematopus ostralegus) and common tern (Sterna hirundo) from Trischen, total PCB concentrations (sum of 32 congeners) declined between 1987
and 2008 from ≈5,500 to < 1,000 ng/g ww (common tern) and from ≈2,500 to < 500 ng/g
ww (oystercatcher) [73] (values derived from figures). Specifically, the data for oystercatcher eggs are
in good agreement with our findings from Trischen. For the Baltic Sea, similar concentrations
were reported in eggs of the little tern (Sterna albifrons) from the east coast of Schleswig-Holstein: Between 1978 and 1996, total PCBs (62
congeners) decreased from ≈3,000 to 600 ng/g ww [74]. Our time series at Heuwiese started later. Nevertheless, in 1996, we observed comparable
∑PCB levels in herring gull eggs, i.e., 820 ng/g ww. In 2008, levels had decreased
to 460 ng/g ww which is in good agreement with the data reported for herring gull
eggs from the Swedish west coast in the same year (sum of 13 congeners 107 to 355
ng/g ww; [75]).

Toxicological implications

OSPAR [76] recommends an ecological quality objective [EcoQO] for total PCBs in eggs of North
Sea seabirds of 20 ng/g ww. The objective was derived for common tern and Eurasian
oystercatcher. It may, however, give an indication of safe levels also in herring
gull eggs as PCB accumulation is quite similar [77]. All of our egg samples from both the Baltic Sea and North Sea exceed the EcoQO of
20 ng/g ww by far. Even single congeners, i.e., CB-118 (with one exception at Mellum
in 2008), CB-138, -153 and -180 exceed the EcoQO at all three sites during the whole
study period. Whether these high levels of PCBs have any effects on herring gull embryos
and hence on the herring gull population is not clear. Studies on effects of PCBs
on herring gull embryos are not available. In laboratory studies with American kestrels
(Falco sparverius), developmental abnormalities and effects on growth were observed at considerably
higher PCB egg concentrations (34 μg/g ww, sum of congeners present in Aroclor 1,248;
1,254; 1,260) [78]. However, besides possible differences in species sensitivities, it is always problematic
to compare laboratory studies with field data. In nature, birds are exposed to a multitude
of contaminants and other environmental stressors, and synergistic and additive effects
may occur. This makes it extremely difficult to assign effects to one pollutant or
stressor only. The only exception may be studies in the vicinity of point sources.
For example, Ormerod et al. [79] studied dippers (Cinclus cinclus) at several unpolluted and one PCB-polluted river section in Wales. They found no
differences in populations even though PCB levels (sum of > 29 congeners) in eggs
at the polluted site were considerably higher (i.e., 8.35 μg/g lw (and thus comparable
to our data) vs. 2.1 μg/g lw at the unpolluted sites). These findings are in line
with other studies that report high pollutant levels in bird eggs but with no associated
population level effects (e.g., [15,80]). Also, for herring gulls at the three ESB sites, no population effects are reported.

PFC

PFOS was the dominant fluorinated compound in herring gull eggs from all three ESB-sampling
sites. PFOA, PFNA, PFDA, PFUnA, PFHxS, and PFHpS were detectable at significantly
lower levels, while PFBA, PFHxA, PFBS, and PFDS concentrations were below the respective
LOQs in most samples.

During the whole study period, mean PFOS levels in herring gull eggs from the North
Sea locations Trischen and Mellum were 91 ± 32 ng/g ww (range 46 to 170 ng/g) and
67 ± 19 ng/g ww (range 39 to 99 ng/g), respectively (Figure 4) [see Table S5 in Additional file 1]. For eggs from Trischen, high concentrations were found throughout the period from
1994 to 2000 as well as in the years 1988, 2002, and 2008 and no temporal trends were
detectable [see Table S7 in Additional file 1]. In contrast, a significant decreasing trend of 3% annually was observed for PFOS
in eggs from Mellum during the entire study period (p < 0.015) and an even stronger trend of 15% annual decrease between the years 2000
and 2006 (p < 0.001) [see Table S7 in Additional file 1]. In 2007 and 2008, levels increased again.

Figure 4.PFOS in herring gull eggs from three ESB locations in German coastal waters. (Mellum and Trischen: North Sea; Heuwiese: Baltic Sea).

At the Baltic Sea location, PFCA egg concentrations were generally lower. Since 1996,
PFUnA and PFDA dominated with mean concentrations of 4 ± 4 ng/g ww and 3 ± 2 ng/g
ww, respectively [see Table S5 in Additional file 1] (Figure 5). With the exception of 2003, concentrations of PFOA were lower (mean 1 ± 1 ng/g
ww; range < LOQ to 3 ng/g). Similar to both North Sea locations, the sums of the concentrations
of longer-chain PFCAs, i.e., PFNA, PFDA, PFUnA, and PFDoA, increased significantly
(p < 0.001) during the study period [see Table S7 in Additional file 1]. This increase may be related to enhanced production volumes of fluorotelomer-based
products (e.g., an approximately two-fold increase between 2000 and 2004), which are
precursors of PFCAs [9,81,82].

The observed differences in PFCA composition between our North Sea sites on one hand
(PFOA dominance) and the Baltic Sea site on the other (more longer-chain PFCAs) might
be related to the feeding habits of the respective gull populations, e.g., more demersal
feeding in the Baltic Sea and thus an increased uptake of particle-associated longer-chain
PFCAs as discussed by Carlsson et al. for eiders [75].

Comparison with other monitoring data

The predominance of PFOS among PFCs in seabird eggs is in accordance with other studies
(e.g., [36,37,40,67,75,83-87]) although concentration levels differ considerably between species and sites.

In guillemot eggs from the Swedish island Stora Karlsö, PFOS increased significantly
from 25 ng/g ww in 1968 to 614 ng/g in 2003. The highest PFOS concentrations were
found in 1997 (1,324 ng/g) and 1999 (1,023 ng/g) [40]. Bignert et al. [88] report mean PFOS concentrations of 1,475 ng/g ww in guillemot eggs from Stora Karlsö
in 2007 and a significant increasing trend of 7% to 10% per year between the late
1960s and the early 2000s. During the last 10 years, however, no further increase
was detectable. Löfstrand et al. [84] compared different northern European locations (i.e., Baltic Sea (Stora Karlsö),
Norway, Faroer Islands, and Iceland) and found the highest PFOS levels in guillemot
eggs from the Baltic Sea with a mean concentration of 400 ng/g ww in 2003. Compared
to Stora Karlsö, eggs from our Baltic Sea sampling site had considerably lower PFOS
levels (20 to 159 ng/g ww).

No comparative data are available for PFOS in bird eggs from North Sea locations.
In herring gull eggs from northern Norway, Verreault et al. [83] detected mean PFOS levels of 21 to 42 ng/g ww in 1983, 1993, and 2003 with a two-fold
increase between 1983 and 1993 and a leveling-off thereafter. Eggs of eiders and shags
from different locations in Norway had mean PFOS levels of 15 to 37 ng/g ww (eider)
and 37 ng/g ww (shag) in 2003 and 2004 [67]. Slightly higher PFOS levels are reported for eggs of ivory gulls from the Russian
and Norwegian Arctic, i.e., 58 to 79 ng/g ww in 2006 [71]. Compared to our data, herring gulls, eiders, and shags from Norway were slightly
less contaminated, whereas ivory gulls of the Arctic were within the same range.

In studies dealing with PFCAs in seabird eggs, typically longer-chain PFCAs (> 8 C-atoms)
predominate, while PFOA is often not detected at all [12,36,71,83-87]. This has been ascribed to a positive correlation between carbon chain length and
bioaccumulation for PFC with 8 to 12 carbons [34].

No comparative PFCA-data are available for bird eggs from the North Sea. However,
for guillemot eggs from Iceland, the Faroer Islands, and Norway [84]; herring gull eggs from northern Norway [83]; and ivory gull eggs from the Russian and Norwegian Arctic [71], a predominance of longer-chain PFCA with > 8 C-atoms was reported although concentration
levels varied strongly.

It needs to be clarified whether differences in feeding habits of the gulls, in contamination
of sites, or in sample treatment are responsible for our contradictory findings in
North Sea samples (i.e., PFOA dominance). With respect to sample treatment, one possible
explanation could be that different extraction procedures were applied. Our protocol
includes both alkaline and acidic treatments and is assumed to characterize the bioavailable
fraction of PFCs. Since acidic conditions are applied, the procedure simulates processes
during food digestion. This may result in higher extraction efficiencies for PFOA
as compared to other methods.

Toxicological implications

Based on dietary exposure studies with mallard ducks (Anas platyrhynchos) and the northern bobwhite quail (Colinus virginatus), Newsted et al. [89] calculated toxicity reference values [TRV] and predicted no-effect concentrations
[PNEC] for PFOS based on characteristics of avian top predators. The conservative
egg yolk-based TRV and PNEC were 1.7 μg PFOS/mL and 1 μg PFOS/mL, respectively (assumption
1 mL egg yolk ≈ 1 g). Compared to these values, the PFOS concentrations found in our
study are much lower and - assuming a similar sensitivity of both species - should
alone not pose a threat to herring gull embryos. For PFCAs, corresponding reference
values are not yet available. Again, it must be kept in mind that PFOS is not the
only contaminant detected in gull eggs and that additive/synergistic effects with
other pollutants are likely to occur. Compared to organochlorines (e.g., PCBs and
DDT) PFCs seem to be less critical for gull populations as Bustness et al. [90] have demonstrated for lesser black-backed gulls (Larus fuscus fuscus) from Norway.

The congener distribution was more or less the same at all three locations with a
clear predominance of BDE-47 and BDE-99 until 2005/2006 [see Table S6 in Additional
file 1]. Roughly, congener patterns can be described as follows:

All other congeners were detected at considerably lower concentrations if at all.
During the monitoring period, BDE-47 and -99 decreased markedly. Declining levels
were also observed for most other congeners except BDE-209.

The major congeners can be allocated to three technical PBDE formulations of commercial
and environmental relevance, i.e., Penta-BDE, Octa-BDE, and Deca-BDE. Congeners BDE-47,
-99, -100, and -153 are the major components of Penta-BDE, whereas Octa-BDE contains
mainly BDE-183 and varying amounts of up to 10 other congeners including BDE-196,
-197, -203, -207, and -209 [91]. Technical Deca-BDE consists of > 97% BDE-209 [91,92]. In data evaluation, this was taken into account by combing the relevant congeners
of Penta-, Octa-, and Deca-BDE, respectively, thus linking concentration trends in
herring gull eggs to commercial products (Figures 6 and 7).

Figure 6.Penta-BDE in herring gull eggs from three ESB locations in German coastal waters. The graph shows the sum of the main Penta-BDE components, i.e., BDE congeners 47,
85, 99, 100, 153, and 154) (Mellum and Trischen: North Sea; Heuwiese: Baltic Sea).

Figure 7.Octa-BDE in herring gull eggs from three ESB locations in German coastal waters. The graph shows the sum of the main Octa-BDE components, i.e., BDE congeners 183,
197, 203, and 207 (Mellum and Trischen: North Sea; Heuwiese: Baltic Sea).

The results clearly show that Penta-BDE was the major contaminant of herring gull
eggs at all three locations even though Deca-BDE production volumes were considerably
higher [43,91]. Penta-BDE concentrations (sum of BDE-47, -85, -99, -100, -153, and -154) ranged
between 89 and 1,615 ng/g lw at both North Sea locations with generally higher levels
in eggs from Mellum (Figure 6). In 1992, however, an exceptionally high Penta-BDE level was detected in eggs from
Trischen which involved all five congeners. It is therefore hypothesized that a single
event, e.g., an accident or spill may be the cause. Penta-BDE levels at the Baltic
Sea location Heuwiese ranged between 176 and 1,128 ng/g lw and were thus comparable
or even slightly higher than levels at Mellum. At all three locations, Penta-BDE decreased
significantly during the study period (p < 0.002 (Mellum), p < 0.005 (Trischen), and p < 0.001 (Heuwiese)) [see Table S7 in Additional file 1]. The declines started already before the EU-wide ban in 2004 but are, nevertheless,
likely to be related to the political discussion (e.g., reduction and substitution
already before the final ban). Fast declines in environmental PBDE concentrations
as a consequence of decreased emissions have also been described by others (e.g.,
[63,93,94]).

Levels of Octa-BDE (sum of BDE-183, 197, 203, and 207) were considerably lower than
those of Penta-BDE. At both North Sea locations, egg concentrations varied between
12 and 124 ng/g lw, again with mostly higher concentrations at Mellum (Figure 7). Similar levels were detected in eggs from the Baltic Sea island Heuwiese (i.e.,
17 and 101 ng/g lw). At all three locations, Octa-BDE increased until 2002 and 2003,
respectively (significant only at Mellum (1988 to 2002), p < 0.001; and Trischen (1988
to 2003), p < 0.002) [see Table S7 in Additional file 1]. Thereafter, concentrations decreased at all sites (significant only at Heuwiese
(p < 0.022, 2003 to 2008) [see Table S7 in Additional file 1]. In 2008, however, Octa-BDE had increased again in eggs from both North Sea locations.
Next years' data will show whether this is indicative of a new rise.

Except for the year 2000, concentrations of Deca-BDE (BDE-209) ranged between 3 and
198 ng/g lw in North Sea eggs with temporarily high inter-year variations. Eggs from
Heuwiese had Deca-BDE levels of 22 to 103 ng/g lw. In 2000, however, unusually high
concentrations of 1,328 ng/g lw (Mellum), 680 ng/g (Trischen), and 995 ng/g (Heuwiese)
were analyzed. The values in year 2000 should be handled with caution as the surprising
high levels have not been confirmed yet. In contrast to Penta-BDE and Octa-BDE, Deca-BDE
increased with time at all three ESB locations; a significant trend, however, was
detected only at Mellum (p < 0.027) [see Table S7 in Additional file 1]. Deca-BDE was restricted in the EU only in 2008, so the following years will show
whether these measures are as effective as the bans of Penta- and Octa-BDE.

Comparison with other monitoring data

In accordance with our findings, congeners BDE-47, -99, and -100 have been described
as dominant congeners in marine bird eggs in numerous studies (e.g., [10,43,65,75,71,94-97]. They are more efficiently bioaccumulated than higher brominated PBDEs [43].

Comparative data for PBDE in marine bird eggs from the Baltic Sea are available for
guillemot eggs from the island Stora Karlsö [65,93] and eider and herring gull eggs from the Swedish west coast [75]. Between 1969 and 2003, BDE-47 in guillemot eggs from Stora Karlsö ranged between
45 and 1,100 ng/g lw, while levels of BDE-99 and -100 were lower (i.e., 2.0 to 190
ng/g and 1.0 to 100 ng/g, respectively, with BDE-100 analyzed only until 1997 [65,93]). Concentrations increased until the end of the 1980s, followed by a rapid decrease
[93]. These values are comparable to the concentrations we detected in herring gull eggs
from our Baltic Sea location. Considerably lower concentrations were detected in eggs
of eiders and herring gulls from the Swedish west coast in 2008 [75]. In both species, BDE-47 was the most abundant congener with mean concentrations
of 0.3 ng/g ww (≈1.6 ng/g lw, calculated with a mean lipid content of 18%) in eider
eggs and 2.6 ng/g ww (≈43 ng/g lw, mean lipid content 6%) in herring gull eggs. In
the same year, 101 ng BDE-47/g lw was detected in our Baltic Sea samples.

Again, no comparative data are available for the North Sea, but several studies have
analyzed PBDEs at different locations in the North Atlantic. For herring gull eggs
from North Norway, Helgason and et al. [94] report rather high BDE-47 concentrations of 350 to 497 ng/g lw in 1983, 1993, and
2003 and considerably lower levels of other congeners. Between 1983 and 1993, levels
increased while they decreased or stabilized thereafter. Lower BDE-47 levels were
detected in eggs of ivory gulls from four locations in the Russian and Norwegian Arctic
(i.e., 25 and 208 ng/g lw in 2006/2007 [71]. Total PBDE concentrations (median of the sum of congeners -28, -47, -99, -100, -153,
-154, -209) at these sites amounted to 51 to 302 ng/g lw. For guillemot from different
North Atlantic locations (Iceland, Faroe Islands, and three Norwegian sites) [65] and fulmars from Faroe Islands [61], mean egg concentrations of 5.9 to 38 ng BDE-47/g lw are reported for years 2000
to 2003 with highest levels in Iceland. Similar BDE-47 concentrations are reported
for eggs of the common eider from three differently exposed locations in Norway (2
to 20 ng/g lw) and eggs of European shags from one of these sites (24 and 27 ng/g
lw) in 2003 and 2004 [67]. For comparison, between years 2000 and 2007, BDE-47 levels in herring gull eggs
from our North Sea locations ranged between 43 and 275 ng/g lw.

As already discussed for PCBs and PFCs, the observed differences between locations
and species are probably related to differences in local contamination as well as
to different feeding habits, metabolism, and ecology. PBDEs are emitted not only by
industrial plants, but also by households, e.g., from electronic equipment, cars,
furniture, carpets, etc., which means that the surroundings of areas with high population
densities are likely to be more contaminated compared to more remote sites [65,67,94]. On the other hand, species-specific properties like feeding habits and metabolism
also influence the contaminant levels in bird eggs [64,67,75] and thus contribute to the reported differences.

Toxicological implications

No reference values for PBDEs are available. Fernie et al. [46,98] report effects on reproduction and behavior of the American kestrel (Falco spaverius) after dietary exposure to two environmental concentrations of the commercial Penta-BDE
formulation DE-71 which resulted in egg concentrations of 288 ± 33 ng/g and 1131 ±
95 ng/g ww (sum of PBDEs), respectively. Following air cell injection, embryonic survival
of American kestrels was affected at absorbed Penta-BDE concentrations of 32 μg/g
egg lw [99]. In herring gull eggs from all three ESB sites, the sum of all 35 investigated PBDE
congeners ranged between 116 and 2,059 ng/g lw (corresponding to 11 to 119 ng/g ww)
during the whole monitoring period and was thus lower than the effect concentrations
for Penta-BDE in American kestrels. Provided that the sensitivity of American kestrels
and herring gulls is comparable, PBDE alone would not lead to effects on reproduction
and embryonic survival. As already stated for PCBs and PFCs, however, herring gulls
are exposed to a broad cocktail of contaminants, and additive/synergistic effects
must be taken into account, e.g., as observed for PCBs and PBDEs [100].

Conclusions

Effects of regulatory measures were clearly observable for PCBs and PBDEs. PCBs in
herring gull eggs have decreased during the study period at all three sites following
the EU-wide ban in 1985. Nevertheless, PCB levels are still high and exceed the common
tern/oystercatcher EcoQOs of 20 ng/g ww by far. However, these high concentrations
are obviously not critical for the herring gulls because population level effects
are not observed at any of the three ESB sites.

In the context of the EU-wide ban of Penta- and Octa-BDE products in 2004 [20], concentrations of the respective congeners, mainly BDE-47, -99, -100 and -183, declined
in herring gull eggs. In contrast, no decreases were observed for BDE-209, the major
component of commercial Deca-BDE which even increased at one site. Since July 2008,
the use of Deca-BDE is restricted within the EU (curia.europa.eu; case C-14/06), and
further monitoring will show whether this is reflected in herring gull eggs.

In contrast to PCBs and PBDEs, the voluntary phase-out in production of PFOS and respective
precursor substances declared by the main manufacturer 3 M in 2000 [101] as well as the restricted use of PFOS in the European Union is currently not mirrored
in decreasing PFOS concentrations in herring gull eggs from two out of three ESB locations.
Decreases in PFOS after 2000 were detectable only in eggs from the North Sea island
Mellum. At both other sites, PFOS concentrations varied between years and even increased
in Heuwiese. Concurrently, an increase of PFCAs with > 8 C-atoms is noticeable in
herring gull eggs since 2000 which may be related to an increased fluorotelomer production
(precursors of PFCAs) compensating the PFOS phase-out. Taking their physicochemical
properties into account, it is likely that PFOS responds quite differently to the
cessation of production compared to lipophilic compounds like PCBs and PBDEs [40]. The coming years will show the effectiveness of the regulatory measures now under
way.

Competing interests

The authors declare that they have no competing interests.

Authors' contributions

AF assessed the data, drafted the manuscript, did the literature research, and performed
the statistical analysis. HR is responsible for the ESB sample preparation and data
compilation and contributed to the drafting of the manuscript. HJ carried out the
chemical analysis of PFCs. JM was responsible for the PFC data compilation. FN was
responsible for the analytical data of PCBs and PBDEs and edited the presentation
of the data. CSK coordinated the research and funding, evaluated the data, and contributed
to the drafting of the manuscript. All authors read and approved the final manuscript.

Acknowledgements

The funding of the German Environmental Specimen Bank project partners by the German
Federal Ministry of the Environment, Nature Conservation and Nuclear Safety is acknowledged.
The authors thank especially the following members of the ESB teams for their valuable
contributions: Sonja Uhlig, Martin Weingärtner (both Fraunhofer IME), Roland Klein,
Markus Quack, Gerhard Wagner (all University of Trier), and Dirk Pfleger (Eurofins
GfA).